The Community Alliance on the Savannah River Site. HARAMBEE HOUSE, INC. Project Citizens For Environmental Justice September 2005

Transcription

1 The Community Alliance on the Savannah River Site HARAMBEE HOUSE, INC. Project Citizens For Environmental Justice September 2005 A Scientific and Technical Review of The Use of Bioremediation as a Clean Up Technique At the Savannah River Site Report Prepared by: Dr. Mildred McClain Scientific and Technical Analysis Conducted by: Michael Ashby, New York City College Dr. Kenneth Sajwan, Savannah State University Dr. Mae Samuel, Benedict College Supported by a grant from the Citizens Monitoring and Technical Assessment Fund

2 -2- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS BIOREMEDIATION

3 -3- ASSESSMENT OF BIOREMEDIATION AS A CLEAN UP STRATEGY AT SAVANNAH RIVER SITE TABLE OF CONTENTS Page I. Summary 3-8 II. Introduction and Background 9 20 III. Bioremediation Technologies IV. Monitored Natural Attenuation and Enhanced Passive Remediation Alternative Project and Technology Acceleration Implementation Plan (DOE) V. The Savannah River Site VI. Monitoring Natural Attenuation VII. References VIII. Glossary and Acronyms

4 -4- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS SUMMARY

5 -5- Halogenated volatile organic compounds (VOCs), including chlorinated aliphatic hydrocarbons (CAHs), are the most frequently occurring type of contaminant in soil and groundwater at Superfund (CERCLA) and other hazardous waste (RCRA) sites in the United States. The U.S. Environmental Protection Agency (EPA) estimates that cleanup of these sites will cost more than $45 billion (1996 dollars) over the next several decades (EPA, 1997). Other evaluations have projected the clean up cost of existing environmental contamination as being more in the region of $1 trillion dollars. In the early 1980's, little was known about how toxic wastes interact with the hydrosphere. This lack of knowledge was crippling efforts to remediate environmental contamination under the new Superfund legislation---the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Faced with this problem, Congress directed the United States Geological Survey (USGS) to conduct a program to provide this critically needed information. By means of this program, known as the Toxic Substances Hydrology Program (TSHP), the most important categories of wastes were systematically investigated at sites throughout the nation. One of the principal findings of this program was that microorganisms in shallow aquifers affect the fate and transport of virtually all kinds of toxic substances. As a result of these findings, innovative technologies, including in situ and ex situ bioremediation, are being developed and implemented in an effort to reduce the cost and time required to clean up currently identified sites. For the most part, these technologies involve biogeochemical processes including the concept known as Natural Attenuation. Natural attenuation may reduce the potential risks posed by site contaminants in one of three ways, depending upon the type of contaminant: 1. Contaminants may be transformed to a less toxic form through destructive processes (e.g., biodegradation, radioactive decay); 2. Potential exposure levels may be reduced by lowering concentration levels (e.g., dilution, dispersion); and 3. Contaminant mobility and bioavailability may be reduced by sorption to the soil or rock matrix. Several organizations involved in environmental clean-up projects have developed definitions for natural attenuation. The U.S. Environmental Protection Agency (EPA) defines monitored natural attenuation as the "reliance on natural attenuation processes (within the context of a carefully controlled and monitored site cleanup approach) to achieve site-specific remediation objectives within a time frame that is reasonable compared to that offered by other more active methods. The 'natural attenuation processes' that are at work in such a remediation approach include a variety of physical, chemical, or biological processes that, under favorable conditions, act without human intervention to reduce the mass, toxicity, mobility, volume, or concentration of contaminants in soil or groundwater. These in-situ processes include biodegradation; dispersion; dilution; sorption; volatilization; radioactive decay; chemical or biological stabilization; transformation; or destruction of contaminants." (EPA, OSWER Directive P) The American Society for Testing and Materials (ASTM) defines natural attenuation as the "reduction in mass or concentration of a compound in groundwater over time or distance from the source of constituents of concern due to naturally occurring physical, chemical, and biological processes, such as; biodegradation, dispersion, dilution, adsorption, and volatilization." (ASTM E , 1998)

6 -6- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS The U.S. Air Force Center for Environmental Excellence defines natural attenuation as the processes resulting "from the integration of several subsurface attenuation mechanisms that are classified as either destructive or nondestructive. Biodegradation is the most important destructive attenuation mechanism. Nondestructive attenuation mechanisms include sorption, dispersion, dilution from recharge, and volatilization." (Wiedemeier, 1999) The U.S. Army defines natural attenuation as "the reduction of contaminant concentrations in the environment through biological processes (aerobic and anaerobic biodegradation, plant and animal uptake), physical phenomena (advection, dispersion, dilution, diffusion, volatilization, sorption/desorption), and chemical reactions (ion exchange, complexation, abiotic transformation). Terms such as intrinsic remediation or bio-transformation are included within the more general natural attenuation definition." (U.S. Army, 1995) Natural attenuation processes for reducing organic contaminant levels are currently best documented at petroleum fuel sites. Organisms in the soil and groundwater break down chemicals through biological degradation processes into byproducts that are often nontoxic and harmless. For example, under appropriate field conditions, the compounds benzene, toluene, ethyl benzene, and xylene (known collectively as BTEX) may naturally degrade through microbial activity and ultimately produce non-toxic end products (e.g., Carbon Dioxide (CO 2 ) and water (H 2 O)). Chlorinated solvents, such as common organic contaminants that may also biodegrade (generally via reductive dechlorination) under certain environmental. Some inorganics, more specifically radionuclides, also break down over time. Unlike organic contaminants, radionuclides have a predictable rate of decay. The specific half-lives of radionuclides allow for accurate prediction of the time required to reduce their radioactivity to levels that are no longer hazardous. The concentrations of mobile and toxic forms of non-degradable inorganic contaminants may also be effectively reduced by other natural processes. The movement of metals and radionuclides is attenuated in the subsurface via sorption to mineral surfaces or soil organic matter and occasionally through volatilization. In addition, oxidation/reduction (redox) reactions can transform the valence states of some inorganic contaminants to less soluble, and thus less mobile, forms, or to forms that are less toxic (e.g., hexavalent to trivalent chromium). Contaminant immobilization through natural processes is contaminant and matrix dependent. Some metals/radionuclides often have very little interaction with the matrix and can, consequently, move unretarded through the subsurface. Furthermore, sorption can be reversible depending upon the contaminant and method of attenuation, i.e., it either becomes a permanent fixture within that particular matrix or maintains the potential for rerelease. Even though some organic and many inorganic contaminants cannot be destroyed or transformed through natural attenuation processes, they are diluted and/or dispersed as they move through the subsurface. Unlike contaminant destruction or transformation, dilution and dispersion do not lead to a reduction in contaminant mass, but rather a reduction in contaminant concentration. Because of the type and occurrence of VOCs at contaminant sites, biodegradation is one of the more important and utilized processes of those mentioned above. What is bioremediation? Biodegradation (biotransformation) is the breakdown of organic contaminants by microbial organisms into smaller compounds. The microbial organisms transform the contaminants through metabolic or enzymatic processes. Biodegradation processes vary greatly, but frequently the final product of the degradation is carbon dioxide or methane and water.

7 -7- Biodegradation is a key process in the natural attenuation of contaminants at hazardous waste sites. The remediation or clean-up of a contaminate site using biodegradation processes and technology is known as Bioremediation. It allows natural processes to clean up harmful chemicals in the environment. Microscopic bugs or microbes that live in soil and groundwater like to eat certain harmful chemicals, such as those found in gasoline and oil spills. When microbes completely digest these chemicals, they change them into water and harmless gases such as carbon dioxide. As technology improves, ex situ clean-ups are becoming less desirable since they require the removal of contaminated materials (soil or groundwater) to be removed to other locations. This can be cost intensive practice. In situ bioremediation is increasingly being selected to remediate hazardous waste sites because, when compared to other technology approaches (e.g. above-ground technologies), it is usually less expensive, does not require waste extraction or excavation, and is more publicly acceptable as it relies on natural processes to treat contaminants. Over half of bioremediation projects at Superfund remedial action sites (57 percent) are in the operational phase, while 26 percent are in the predesign, design, or installation phases and 17 percent have been completed. Of the 18 completed projects, 14 are ex situ source treatment projects, and 4 are in situ projects for source treatment and groundwater treatment. Since 1991, the percentage of bioremediation projects performed ex situ has decreased while the percentage of projects performed in situ has increased. In 1991, only 35 percent of the Superfund remedial action bioremediation projects were in situ versus 53 percent in Bioventing is the most commonly implemented in situ treatment technology for source treatment. Land treatment is the most commonly used ex situ source treatment technology (EPA, 2001; FRTR, 2001). After the initial clean-up of a particular site, it may become necessary to monitor the behavior of the contaminant (plume). This extended oversight has become known as Monitored Natural Attenuation (MNA). Monitored natural attenuation may be defined as the reliance on natural attenuation processes, within the context of a carefully controlled and monitored site cleanup, to achieve site-specific remedial objectives within a time frame that is reasonable compared to that offered by more active methods. Monitoring, therefore, is the critical component of any remediation by natural attenuation. Monitoring is imperative to: o ensure performance objectives are being achieved as expected and o detect unacceptable migration of contamination so that contingency measures can be implemented to prevent any unacceptable risks to human health and the environment. How does bioremediation work? In order for microbes to clean up harmful chemicals, the right temperature, nutrients (fertilizers), and amount of oxygen must be present in the soil and groundwater. These conditions allow the microbes to grow and multiply and eat more chemicals. When conditions are not right, microbes grow too slowly or die. Or they can create more harmful chemicals. If conditions are not right at a site, EPA works to improve them. One way they improve conditions is to pump air, nutrients, or other substances (such as molasses) underground. Sometimes microbes are added if enough aren t already there. The right conditions for bioremediation cannot always be achieved underground. At some sites, the weather is too cold or the soil is too dense. At such sites, EPA might dig up the soil to clean it above ground where heaters and soil mixing help improve conditions. After the soil is dug up, the proper nutrients are added. Oxygen also may be added by stirring the mixture or by

8 -8- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS forcing air through it. However, some microbes work better without oxygen. With the right temperature and amount of oxygen and nutrients, microbes can do their work to bioremediate the harmful chemicals. Sometimes mixing soil can cause harmful chemicals to evaporate before the microbes can eat them. To prevent these chemicals from polluting the air, EPA mixes the soil inside a special tank or building where chemicals that evaporate can be collected and treated. Microbes can help clean polluted groundwater as well as soil. To do this, EPA drills wells and pumps some of the groundwater into tanks. Here, the water is mixed with nutrients and air before it is pumped back into the ground. The added nutrients and air help the microbes bioremediate the groundwater. Groundwater can also be mixed underground by pumping nutrients and air into the wells. Once harmful chemicals are cleaned up and microbes have eaten their available food, the microbes die. Is bioremediation safe? Yes. Bioremediation is very safe because it relies on microbes that naturally occur in soil. These microbes are helpful and pose no threat to people at the site or in the community. While the Microbes themselves won t hurt you, it is always good practice never touch polluted soil or groundwater especially before eating. No dangerous chemicals are used in bioremediation. The nutrients added to make the microbes grow are fertilizers commonly used on lawns and gardens. Because bioremediation changes the harmful chemicals into water and harmless gases, the harmful chemicals are completely destroyed. Why use bioremediation? The EPA estimates that, over the next several decades, site owners will spend billions of dollars to cleanup these sites. New technologies that are less costly and more effective are needed to accomplish hazardous waste site remediation. The selection of bioremediation a cleanup remedy is desirable for several reasons. First and most importantly, once chemical contaminants get into the subsurface, it is virtually impossible to clean them up to pre contamination conditions. Even if there is an attempt to remove all contaminated media. Secondly, bioremediation or some kind of nature based sustainable process is the only option for cleanup. Bioremediation involves harnessing natural processes. At some sites, natural microbial processes can remove or contain contaminants without human intervention. In these cases where natural attenuation (intrinsic bioremediation) is appropriate, substantial cost savings can be realized. Remediation and Clean-up organizations use intrinsic bioremediation because it takes advantage of natural processes. Thirdly, it facilitates the treating contamination in place. Most of the cost associated with traditional cleanup technologies is associated with physically removing and disposing of contaminated soils. Because engineered bioremediation can be carried out in place by delivering nutrients to contaminated soils, it does not incur removal-disposal costs. Polluted soil and groundwater can be cleaned at the site without having to move them somewhere else. If the right conditions exist or can be created underground, soil and groundwater can be cleaned without having to dig or pump it up at all. This allows cleanup workers to avoid contact with polluted soil and groundwater. It also prevents the release of harmful gases into the air. Because microbes change the harmful chemicals into water and harmless gases, few if any wastes are created.

9 -9- Fourthly, it reduces environmental stress because bioremediation methods minimize site disturbance compared with conventional cleanup technologies, post-cleanup costs can be substantially reduced. Scientific objectives and technical approaches are based on the hypothesis that among others the most important mass-removal process for natural attenuation is biodegradation. In addition, it can be further hypothesize that a common pattern of biodegradation activity can be found in most groundwater pollution plumes. There are several different types of computer programs and models that can assist decision makers, regulators, and project managers at contaminated sites and their advisors to be able to forecast natural attenuation for a wide range of pollutants under a variety of site conditions. This will give them, and regulators, greater confidence in including natural attenuation as a viable alternative in their risk-based strategy for site remediation and subsequent management. This will allow resources to be focused on more important environmental and economic objectives at the national levels. Once identify, these zones will have better conditions for biodegradation, and these zones will have more rapid degradation and make a significant contribution to the overall rate of mass loss for the entire plume. Since these areas, once identified, can be modeled prior to actual start up of the clean-up process, there is greater confidence in forecasting the outcome of Natural Attenuation as a risk-based soil or groundwater remediation strategy. Finally, it reduces clean-up costs by stretching remediation dollars. It has been projected that the cost of remediating contaminated sites to meet the minimal regulatory requirements is likely to run into many billions of dollars. For many States and local communities, remediation of contaminated sites may not be economically or technically feasible with conventional cleanup technologies. Information about the cost of using bioremediation to treat contaminated media was available for 67 sites. Unit costs for bioventing projects ranged from approximately $2 per cubic yard (cy) to more than $300/cy, with most sites less than $40/cy. Unit costs for ex situ bioremediation of soil, such as land treatment or composting systems, ranged from $13/cy to more than $500/cy, with most projects costing less than $300/cy. Simply stated, in situ bioremediation can help contain costs (EPA, 2001; FRTR, 2001). As an alternative, a greater emphasis can be placed on risk-based approaches to soil and groundwater resource management. Where remediation objectives at particular contaminated sites are balanced against expected cleanup costs and intended use of the groundwater or land area. Often bioremediation does not require as much equipment or labor as most other methods. Therefore, it is usually cheaper. Few bioremediation Records of Decision (RODs) were signed in the early- to mid-1980s. Beginning in fiscal year (FY) 1988, the number of bioremediation RODs has increased. In general, 8 to 12 bioremediation RODs have been signed per year (EPA, 2001; FRTR, 2001). How long will bioremediation take? The time it takes to bioremediate a site varies depending on several factors: o types and amounts of harmful chemicals present o size and depth of the polluted area o type of soil and the conditions present o whether cleanup occurs above ground or underground These factors vary from site to site. It can take a few months or even several years for microbes to eat enough of the harmful chemicals to clean up the site. To ensure that bioremediation is working, project managers periodically take and analyze samples of soil and groundwater from the contaminated area.

10 -10- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS While Monitored natural attenuation (MNA) has shown that it is a potentially low-cost valuable risk-based remediation strategy for contaminated groundwater, its wider exploitation throughout the nation is still limited by lack of confidence in its application and management, and uncertainty in predicting its performance at many sites.

11 Introduction and Background -11- INTRODUCTION AND BACKGROUND

12 -12- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS This report provides an overview of the fundamentals and field applications of in situ bioremediation in contaminated soil and groundwater. Bioremediation has been presented to the impacted communities as a technology that is currently being used and tested at the Savannah River Site (SRS). The primary focus this report explain bioremediation as a clean-up technology and analyze what role the technology will play in the new accelerated clean up strategy at SRS as approved by Environmental Management at the U.S. Department of Energy (DOE). The objective is to understand how bioremediation is being used at SRS for clean up. Included is a summary of currently-available information on the mechanisms and technologies used to implement in situ bioremediation. Under NEPA (1969), Executive Order and several EPA statues cleanup managers are required to involve the public so that they can become aware of and participate in activities related to cleanup and environmental management ongoing in their own communities. This report is intended to familiarize affected communities and others involved with hazardous waste site cleanups, including site project managers, contractors, and other technology users, with the dynamics of in situ bioremediation. In addition, the report is intended to present a discussion to community residents and members of the working group to help them to identify and develop the knowledge base of the surrounding community on the Performance Management Plan (PMP) and bioremediation as a remedial technology at SRS. As such, the level of detail included in this report about bioremediation mechanisms, technologies, and implementation is meant to provide only basic information about the technology, rather than providing an in depth how to manual about in situ bioremediation. It will be part of a comprehensive picture of the accelerated clean up plan and the use bioremediation as a clean up technology. The report looks at the use and viability of bioremediation in other locations as part of this review. The report should therefore be used for informational purposes, and should not be used as the sole basis for determining the value and usefulness of this technology at the specific site of concern. It is instructive that such decisions must be made on a case-by-case basis, considering site-specific factors. This introduction provides background information about in situ bioremediation at sites contaminated with various examples of chlorinated aliphatic hydrocarbons (CAHs). This type of contamination is chosen because of the type, number of occurrences and difficulty of clean-up. Because of their nature several of these can be further classified as Persistent Organic Pollutants (POPs). The basic description below includes; identification of typical CAHs and their physical and chemical properties; Processes that transport CAHs through the subsurface environment and biological and chemical mechanisms that can degrade CAHs. More detailed information about the physical and chemical characteristics and the subsurface transport processes of CAHs are beyond the scope of this report but can be found in the references listed below. BIOREMEDIATION MECHANISMS Bioremediation has successfully cleaned up many polluted sites and is being used at 50 Superfund sites across the country. The most common type of Superfund remedial action site where bioremediation is used is wood preserving (31 percent), followed by petroleum sites (21 percent). The most common types of contaminants at these sites are polycyclic aromatic hydrocarbons (PAHs) (40 percent); benzene, toluene, ethylbenzene, and xylenes (BTEX) (37 percent); and pesticides and herbicides (27 percent). Available performance data shows that bioremediation is capable of reducing contaminant concentrations in contaminated media. Bioremediation is being used to treat recalcitrant organic compounds, including chlorinated volatile organic compounds (VOCs), PAHs, pesticides and herbicides, and explosives.

13 Introduction and Background -13- For ten projects treating chlorinated VOCs, concentrations of VOCs in treated groundwater ranged from below detection limit (<5 µg/l for tetrachloroethene [PCE], trichloroethene [TCE], and dichloroethene [DCE]) to 1,200 µg/l (for carbon tetrachloride). For seven projects treating PAHs, concentrations of PAHs in treated soil and sludges ranged from 3.3 mg/kg to 795 mg/kg, with some projects showing more than 90% removal. For four projects treating pesticides and herbicides, concentrations of specific pesticides and herbicides in treated soil were less than 10 mg/kg at two projects and less than 200 mg/kg at the other two projects, with some projects showing more than 90% removal. For six projects treating explosives, three showed removals of more than 75% and the others showed removals ranging from little or none to as much as 64% (EPA, 2001; FRTR, 2001). Chlorinated aliphatic hydrocarbons (CAHs) are manmade organic compounds. They are typically manufactured from naturally occurring hydrocarbon constituents (methane, ethane, and ethene) and chlorine through various processes that substitute one or more hydrogen atoms with a chlorine atom, or selectively dechlorinate chlorinated compounds to a less chlorinated state. As such CAHs are also classified as chlorinated volatile organic compounds (VOCs). They are used in a wide variety of applications, including use as solvents and degreasers and in the manufacturing of raw materials. They include such solvents as tetrachloroethene (PCE), trichloroethene (TCE), carbon tetrachloride (CCl4), chloroform (CF), and methylene chloride (MC). Historical management of wastes containing CAHs has resulted in contamination of soil and groundwater, with CAHs present at many contaminated groundwater sites in the United States. Of these substances, trichloroethene (TCE) is the most prevalent of those contaminants (U.S. Air Force, 1998). In addition to CAHs, their degradation products can include such substances as dichloroethane (DCA), dichloroethene (DCE), and vinyl chloride (VC), which tend to persist in the subsurface. Physical and Chemical Properties of CAHs The physical and chemical properties of CAHs govern their transport and fate in the subsurface (underground) environment. The number of substituted chlorine atoms on the CAHs directly affects their physical and chemical behavior. As the number of substituted chlorine atoms increases, the molecular weight and density generally increase, whereas vapor pressure and aqueous solubility generally decrease. A CAH released to the subsurface as a pure organic liquid (commonly referred to as non-aqueous phase liquid [NAPL] in the subsurface) will seek phase equilibrium (a condition in which all acting influences are canceled by others, resulting in a stable, balanced, or unchanging system). The CAH will remain as a NAPL, adsorb to soil, dissolve in groundwater, or volatilize into soil gas to the extent defined by the physical and chemical properties of the individual CAH and the subsurface environment. Partition coefficients, which are related to the hydrophobicity and aqueous solubility of a CAH, define the extent to which a CAH will partition between NAPL, adsorb to soil, and dissolve in groundwater. The vapor pressure of a CAH defines the extent to which it will partition between NAPL or NAPL adsorbed to soil and the soil gas. CAHs dissolved in groundwater will also partition themselves between the dissolved phase and the vapor phase, as defined by their Henry s Constant. Figure 1 below shows those mechanisms by which CAHs transfer phases in an attempt to reach equilibrium conditions and their related properties. Non-Aqueous Phase Liquids (NAPLs) are liquids that are sparingly soluble in water. Because they do not mix with water, they form a separate phase. Hydrocarbons, such as oil and

14 -14- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS gasoline, and chlorinated solvents, such as trichloroethylene, are examples of NAPLs, because they do not mix with water, and oil and water in a glass will separate into two separate phases. Figure 1: Phase Equilibrium Mechanisms and Defining Properties of CAHs Source: Modified from Huling and Weaver, 1991 There are two categories of NAPLs. The first type can be lighter than water (LNAPL) while the second group is denser than water (DNAPL). Most of the CAH NAPLs discussed in this report are denser than water. They are generally referred to as dense non-aqueous phase liquids [DNAPLs]). The exceptions are vinyl chloride, chloroethane, and chloromethane, which are gaseous in their pure phase under standard conditions. NAPLs that are less dense than water are generally referred to as light non-aqueous phase liquids [LNAPL]. In addition, capillary forces can trap NAPLs in porous media above or below the water table. Light Non-Aqueous Phase Liquids (LNAPLs) are liquids that are sparingly soluble in water and less dense than water. They will sink through unsaturated permeable soils and float on the water table, migrating to the lowest water table elevation. Hydrocarbons, such as oil and gasoline, are examples of LNAPLs, oil will "float" on top of water and does not mix. At LNAPL contamination sites, LNAPL can form a pool in the subsurface on top of the water table. The following diagram is a cross sectional view of a hypothetical LNAPL spill. Dense Non-Aqueous Phase Liquids (DNAPLs) are liquids that are denser than water and do not dissolve or mix easily in water (they are immiscible). DNAPLs will tend to sink through both unsaturated (vadose) and saturated (phreatic) zones of permeable soils until they reach the lowest point on the top of a confining layer of the aquifer or bedrock. This movement is described by some as immiscible transport. In the presence of water it forms a separate phase from the water. Many chlorinated solvents, such as trichloroethylene are DNAPLs. TRANSPORT PROCESSES

15 Introduction and Background -15- Figure 2: Typically generalized contaminant flow model showing plume dynamics In addition to transferring phases in an attempt to reach equilibrium conditions (fig. 1), CAHs can migrate in the subsurface in their non-aqueous, aqueous, and vapor phases by both active and passive processes. The active processes, involves such dynamics as advection, dispersion, sorption and relative mobility. CAHs migrate along with the flow of the groundwater or soil gas to which they are partitioned. Passive processes generally involve diffusion and are the result of concentration gradients, which cause the CAH to seek phase and concentration equilibrium with its surrounding environment. The extent of subsurface migration is a function of the volume of CAH released; the area over which the release occurs; the duration of the release; the chemical and physical properties of both the CAH and the subsurface environment. Typically, releases of CAHs to the groundwater result in the formation of a plume (see figure 2 above); releases to soil result in subsurface soil contaminated with CAH constituents. In soil, CAHs typically are transported by the flow of DNAPL or diffusion in soil-gas vapor. In groundwater, advective transport (the movement of contaminants by flowing groundwater) is one of the most important processes that affect the transport of dissolved CAHs. In general, the more soluble the compound, the further it will be transported in the subsurface. For example, based on solubility data provided in figure, MC and CF would be transported more readily in groundwater that PCE and CT. Figure 3 presents a more detailed example of the subsurface transport processes associated with the dense non-aqueous phase liquid [DNAPL] trichloroethene (TCE), a very prevalent component of those subsurface contamination.

16 -16- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS Figure 3: Example CAH Subsurface Transport Processes (DNAPL Source) Source: Modified from Sims et al., 1992 DEGRADATION MECHANISMS Bioremediation of CAHs can occur through natural mechanisms (intrinsic bioremediation) or by enhancing the natural mechanisms (enhanced bioremediation). For a few CAHs (for example, 1,1,1-TCA and CCl4), degradation can also occur by abiotic (nonbiological) mechanisms. CAHs can also be degraded or otherwise removed from soil and groundwater by larger organisms (such as trees), in a process referred to as phytoremediation. In most systems, biological degradation tends to dominate, depending on the type of contaminant and the groundwater chemistry (EPA, 1998). A number of biological degradation mechanisms have been identified theoretically and observed on a laboratory scale. The bioremediation mechanisms carried out by bacteria that typically are used for enhanced bioremediation of CAHs generally can be classified into one of the following mechanism two categories: - Aerobic oxidation (direct and cometabolic) - Anaerobic reductive dechlorination (direct and cometabolic) While aerobic oxidation and anaerobic reductive dechlorination can occur naturally under the proper conditions, enhancements such as the addition of electron donors, electron acceptors, or nutrients can help to provide the proper conditions for aerobic oxidation or anaerobic reductive dechlorination to occur. In general, highly chlorinated CAHs degrade primarily through reductive reactions, while less chlorinated compounds degrade primarily through oxidation (Vogel et al., 1987b). Highly chlorinated CAHs are reduced relatively easily because their carbon atoms are highly oxidized. During direct reactions, the microorganism causing the reaction gains energy or grows as the CAH is degraded or oxidized. During cometabolic reactions, the CAH degradation or oxidation is caused by an enzyme or cofactor produced during microbial metabolism of another compound. CAH degradation or oxidation does not yield any energy or growth benefit for the microorganism mediating the cometabolic reaction. Biodegradation involves the production of energy in a redox reaction within a bacterial system.

17 Introduction and Background -17- This includes respiration and other biological functions needed for cell maintenance and reproduction. Ecology involves the different types of bacteria electron acceptor classes, such as oxygen-, nitrate-, manganese-, iron (III)-, sulfate-, or carbon dioxide-reducing, and their corresponding redox potentials. Redox potentials provide an indication of the relative dominance of the electron acceptor classes. Aerobic Oxidation In aerobic zones (zones of the subsurface where oxygen is present) of the subsurface, certain CAHs can be oxidized to carbon dioxide, water, and chloride by direct and cometabolic mechanisms (Hartman and DeBont, 1992; McCarty and Semprini, 1994; Malachowsky et al., 1994; Gerritse et al., 1995; Bielefeld et al., 1995; Hopkins and McCarty, 1995). Direct mechanisms are more likely to occur with the less chlorinated CAHs (mono- and di-chlorinates). In general, the more chlorinated CAHs can be oxidized by cometabolic mechanisms, but no energy is provided to the organism. Incidental oxidation is caused by enzymes intended to carry out other metabolic functions. Generally, direct oxidation mechanisms degrade CAHs more rapidly than cometabolic mechanisms (McCarty and Semprini, 1994) Figure 4: aerobic oxidation (direct) and (cometabolism) of CAH Aerobic Oxidation (Direct) Aerobic oxidation (direct) is the microbial breakdown of a compound in which the compound serves as an electron donor and as a primary growth substrate for the micro organism (microbe) mediating the reaction. Electrons that are generated by the oxidation of the compound are transferred to an electron acceptor such as oxygen. In addition, a microorganism can obtain energy for cell maintenance and growth from the oxidized compound (the compound acts as the reductant). In general, only the less chlorinated CAHs (CAHs with one or two chlorines) can be used directly by microorganisms as electron donors. CAHs that can be oxidized directly under aerobic conditions include DCE, DCA, VC, CA, MC, and CM (RTDF, 1997; Bradley, 1998; Harkness et al., 1999). The CAHs are oxidized into carbon dioxide, water, chlorine, and electrons, in conjunction with the reduction of oxygen to water. Figure 4 shows an example of aerobic oxidation (direct) of a CAH. Aerobic Oxidation (Cometabolic) Aerobic oxidation (cometabolic) is the microbial breakdown of a contaminant in which the contaminant is oxidized incidentally by an enzyme or cofactor produced during microbial

18 -18- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS metabolism of another compound. In such a case, the oxidation of the contaminant does not yield any energy or growth benefit for the microorganism involved in the reaction. The CAHs that have been observed to be oxidized cometabolically under aerobic conditions include TCE, DCE, VC, TCA, DCA, CF, and MC (Edwards and Cox, 1997; McCarty, 1997a; Munakata-Marr, 1997; RTDF, 1997; Travis and Rosenberg, 1997; Bradley and Chapelle, 1998; McCarty et al., 1998). The electron donors observed in aerobic oxidation (cometabolic) include methane, ethane, ethene, propane, butane, aromatic hydrocarbons (such as toluene and phenol), and ammonia. Under aerobic conditions, a monooxygenase (methane monooxygenase in the case of methanotrophic bacteria) enzyme mediates the electron donation reaction. That reaction has the tendency to convert CAHs into unstable epoxides (Anderson and Lovley, 1997). Unstable epoxides degrade rapidly in water to alcohols and fatty acids, which are readily degradable. Figure 4 above also shows an example of aerobic oxidation (cometabolic) of a CAH. Wilson and Wilson (1985) were the first to observe that the simultaneous addition of methane and oxygen can stimulate biodegradation by aerobic oxidation (cometabolic) of TCE in aquifer material. Subsequently, that approach was tested in the field at Naval Air Station (NAS) Moffett Field, California. Intermittent pulses of oxygen and methane were provided to the subsurface, bringing about the in situ stimulation of biodegradation of TCE, c-dce, and VC in a contaminated aquifer (Semprini et al., 1990). The strategy has been applied successfully to biodegradation of CAHs at a variety of other sites (McCarty et al., 1991; Travis and Rosenberg, 1997). Although the studies have demonstrated that addition of methane is an effective means of stimulating cometabolic biodegradation of CAHs, additional field studies at the Moffett test site have shown that toluene and phenol can be more effective electron donors than methane in the stimulation of cometabolic biodegradation of TCE, c-dce, and VC in groundwater (Hopkins et al., 1993; Hopkins and McCarty, 1995). Anaerobic Reductive Dechlorination Under anaerobic conditions, reductive dechlorination mechanisms can effectively biodegrade CAHs. Reductive dechlorination generally involves the sequential replacement of a chlorine atom on a CAH with a hydrogen atom (that is, converting PCE to TCE to DCE, and so on) and has been observed to occur both directly and co-metabolically. In anaerobic reductive dechlorination (direct), the mediating bacteria use the CAH directly as an electron acceptor in energy-producing redox reactions. Anaerobic reductive dechlorination (cometabolic) occurs when bacteria incidentally dechlorinate a CAH in the process of using another electron acceptor to generate energy. Reductive dechlorination theoretically is expected to occur under most anaerobic conditions, but has been observed to be most effective under sulfate-reducing and methanogenic conditions (EPA 1998). As in the case of aerobic oxidation, the direct mechanisms may biodegrade CAHs faster than co-metabolic mechanisms (McCarty and Semprini, 1994) Anaerobic Reductive Dechlorination (Direct) Anaerobic reductive dechlorination (direct) is a biodegradation reaction in which bacteria gain energy and grow as one or more chlorine atoms on a chlorinated hydrocarbon are replaced with hydrogen (Fennel et al., 1997; McCarty, 1997b; Mayo-Gatell et al., 1997; Gerritse et al., 1999). In that reaction, the chlorinated compound serves as the electron acceptor, and hydrogen serves as the direct electron donor (Fennel et al., 1997). Hydrogen used in the reaction typically is supplied indirectly through the fermentation of organic substrates. The reaction is also referred to as halorespiration or dehalorespiration (Gossett and Zinder, 1997).

19 Introduction and Background -19- Anaerobic reductive dechlorination (direct) has been observed in anaerobic systems in which PCE, TCE, DCE, VC, and DCA are used directly by a microorganism as an electron acceptor in their energy producing redox reactions (Freedman and Gossett, 1989; Major et al., 1991; DeBruin et al., 1992; Hollinger, 1993; Hollinger and Schumacher, 1994; Neumann et al., 1994; Tandol, 1994; Yagi et al., 1994; Scholz-Muramatsu et al., 1995; Gerritse et al., 1996; Gossett and Zinder, 1996; Sharma and McCarty, 1996; Smatlak, 1996; McCarty, 1997b; Maymo-Gatell et al., 1997; Yager et al., 1997). The mechanism generally results in the sequential reduction of a chlorinated ethene or chlorinated ethane to ethene or ethane. Figure 5 shows the step-by-step dechlorination of PCE. The anaerobic reductive dechlorination of the more chlorinated CAHs (PCE and TCE) occurs more readily than the dechlorination of CAHs that already are somewhat reduced (DCE and VC); for that reason, DCE and VC may accumulate in anaerobic environments. It also has been observed that, while VC can be effectively dechlorinated, the presence of PCE in groundwater may inhibit the anaerobic reductive dechlorination of VC (Tandol and others, 1994). VC is more commonly remediated using aerobic mechanisms than anaerobic mechanisms. In anaerobic environments in which VC accumulates, enhanced aerobic bioremediation can be implemented to degrade the VC. Recent studies have demonstrated significant anaerobic oxidation of VC to carbon dioxide under Fe(III)-reducing conditions (Bradley and Chapelle 1998b) and of DCE to VC and VC to carbon dioxide under humic acidreducing conditions (Bradley and Chapelle 1998a). These studies suggest the possibility of alternative biotransformation mechanisms under anaerobic conditions. Figure 5: diagram shows the step-by-step dechlorination of PCE. Hydrogen has been observed to be an important electron donor in anaerobic reductive dechlorination (Fennell et al., 1997). The presence of hydrogen establishes a competition between the bacteria that mediate the anaerobic reductive dechlorination (such as Dehalococcus ethenogenes and Dehalospirillium multivorans) and methanogenic bacteria that also use hydrogen as an electron donor (ITRC 2000). However, it has been observed that the dechlorinating bacteria can survive at a partial pressure of hydrogen ten times lower than that at which the methanogenic bacteria can survive (Smatlak et al., 1996), thus providing an opportunity to support the dechlorinating bacteria by providing hydrogen at a slow rate. Hydrogen addition at a slow rate has been demonstrated with the fermentation of butyric or propanoic acid (Fennell et al., 1997). In addition, in some subsurface environments, competition from nitrate or sulfate-reducing bacteria may limit both methanogenic activity and the extent of anaerobic reductive dechlorination (RTDF, 1997).

20 -20- Community Alliance on the Savannah River Site (SRS) Assessment of Bioremediation as a Clean-up Strategy at SRS Past studies have shown that anaerobic reduction of CAHs can occur by reductive dechlorination in a variety of environmental conditions (Beeman et al., 1994; Semprini et al., 1995). A review of the transformation of halogenated compounds has shown that the theoretical maximum redox potential for transformation of PCE to TCE is +580 millivolts and for TCE to DCE is +490 millivolts (Vogel et al., 1987). Therefore, the anaerobic reductive dechlorination of the compounds is thermodynamically possible under manganese- or iron-reducing conditions. No peer-reviewed reports of the transformation of PCE to TCE under aerobic conditions were identified. However, the efficiency of the anaerobic dechlorination processes at high redox potential values is limited; efficiency improves as the redox potential decreases. Figure 6 shows the possible reduction/oxidation (REDOX) zones associated with a petroleum plume in an aerobic aquifer. Pilot studies have been conducted at a variety of sites to examine the feasibility of stimulating in situ anaerobic reductive dechlorination by providing to the subsurface simple organic substrates, such as lactate, butyrate, methanol, ethanol, and benzoate (Freedman and Gossett, 1989; Gibson and Sewell, 1992; Becvar et al., 1997; Buchanan et al., 1997; Litherland and Anderson, 1997; Spuij et al., 1997; Sewell et al., 1998; Harkness et al., 1999). Figure 6: Redox Zones of a Typical Petroleum Plume in an Aerobic Aquifer (Areal View) Source: Modified from Anderson and Lovley, 1997 Anaerobic Reductive Dechlorination (Cometabolic) Anaerobic reductive dechlorination (cometabolic) is a biodegradation reaction in which a chlorinated hydrocarbon is fortuitously degraded by an enzyme or cofactor produced during microbial metabolism of another compound. In such a case, biodegradation of the chlorinated compound does not appear to yield any energy or growth benefit for the microorganism mediating the reaction (Gossett and Zinder, 1997). Several CAHs have been observed to be reductively dechlorinated by co-metabolic mechanisms. In those instances, the enzymes that are intended to mediate the electron-accepting reaction accidentally reduce and dehalogenate the CAH. Anaerobic reductive dechlorination (co-metabolic) has been observed for PCE, TCE, DCE, VC, DCA, and CT under anaerobic conditions (Fathepure, 1987; Workman, 1997; Yager et al., 1997). Combined Aerobic Oxidation and Anaerobic Reductive Dechlorination

21 Introduction and Background -21- Several investigators have suggested that the most efficient bioremediation of CAHs will occur in aquifers that are characterized by an upgradient anaerobic zone and a downgradient aerobic zone (Fathepure et al., 1987; Carter and Jewell, 1993; Bouwer, 1994; Gerritse et al., 1995). In the upgradient aerobic zone, anaerobic reductive dechlorination of PCE might degrade to TCE, and eventually to VC. VC could then be degraded by aerobic oxidation (direct) downgradient in the aerobic zone of the CAH plume (the leading-edge fringe of the plume). Stratified redox conditions in the field may provide the best opportunities, other than engineered remedies, for intrinsic biodegradation of CAHs. Generally, the substrate requirement for direct metabolism is relatively less than that for co-metabolism. In co-metabolism, often the amount of primary substrate required is a factor of 100 to 1,000 times the amount of the CAH. In direct metabolism (respiration with only the chlorinated solvent as the electron acceptor), the stoichiometry is much more favorable, and a much smaller amount of supplemental chemical is required (Bouwer, 1994). Abiotic Degradation Mechanisms Abiotic degradation mechanisms involve chemical reactions to treat CAHs without biological processes. These mechanisms include hydrolysis, elimination, and abiotic reductive dechlorination. In general, the rates of abiotic degradation may be slow relative to biological mechanisms. However, the abiotic mechanisms may play a significant role in the overall remediation of a site at which CAH contamination is present, depending on the specific site conditions (for example, a site at which the contaminant plume is moving slowly) (EPA, 1998). Hydrolysis and elimination reactions are generally independent of redox conditions, while abiotic reductive dechlorination is highly dependent on redox conditions. Hydrolysis is a substitution reaction in which a CAH may react with water to substitute a chlorine atom with a hydroxyl group, producing organic alcohols, acids, or diols, such as the formation of acetic acid from 1,1,1-TCA equation (1). Generally, less chlorinated CAHs are more susceptible to degradation by hydrolysis. Hydrolysis rates have been reported that have half-lives ranging from days for monochlorinated alkanes to thousands of years for tetrachloromethane. Cl 3 C-CH 3 + 2H 2 O H 3 C-COOH + 3HCl (1) 1,1,1-TCA Acetic Acid Hydrolysis is a common transformation mechanism for 1,1,1-TCA, chloroethane, and chloromethane, producing acetate, ethanol, and methanol, respectively (Vogel and McCarty, 1987). Elimination reactions involve the removal of a hydrogen and a chlorine atom (sometimes referred to as dehydrohalogenation) from a chlorinated alkane, with the formation of the corresponding alkane equation (2). In contrast to hydrolysis reactions, elimination reactions become more effective as the CAHs become more chlorinated. Assuming that elimination rates for monochlorinated CAHs are negligible, the abiotic conversion of TCA to DCE at 20oC has been reported to exhibit relatively rapid first-order kinetics, with a rate constant of approximately 0.04 ± year-1 (Vogel and McCarty, 1987). Cl 3 C-CH 3 Cl 2 = CH 2 + HCl (2) 1,1,1-TCA 1,1-DCE